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332 marcel rejmánek ET AL.13

Plant invasions and invasibilityof plant communities

Marcel Rejmánek, David M. Richardsonand Petr Py1ek

13.1 Introduction

Some 2500 yr ago, Heraclitus of Ephesus said that ‘All things change . . . and youcannot step twice into the same stream’. Today, ecologists would not only say thesame about streams but also about vegetation. Plant communities change with timedue to changes in the environment (Chapter 7), biotic interactions (Chapter 9) andinvasions of alien species and genotypes, introduced intentionally or accidentally byhumans. Invasions have received detailed attention only recently. There have alwaysbeen migrating taxa, but now the rate of human-assisted introductions of new taxa isseveral orders of magnitude higher. In California, for example, more than 1000 alienplant species were intentionally or non-intentionally introduced and established viablepopulations over the last 250 yr. In the Galápagos Islands, over 3 million yr of theirhistory, only one new plant species arrived with birds or sea currents every 10,000 yr.However, over the last 20 yr the introduction rate has been c. 10 species per yr, orsome 100,000 times the natural arrival rate (Tye 2001).

Three basic questions arise:1 What kind of ecosystems are more (or less) likely to be invaded by alien plants?2 What kind of plants are the most successful invaders and under what circumstances?3 What is the impact of the plant invaders?

13.2 Definitions and major patterns

Unlike natives (taxa that evolved in the region or reached it from another area wherethey are native without help from humans), aliens (‘non-native’ or ‘exotic’) owe theirpresence to direct or indirect activities of humans. Most of them occur only tem-porarily and are not able to persist for a long time without human-assisted input ofdiaspores; these are termed casual. Naturalized taxa form sustainable populationswithout direct human help but do not necessarily spread; the ability to spread char-acterizes a subset termed invasive taxa. This distinction is critical because not allnaturalized taxa reported in floras and checklists are invasive. Not all naturalizedplant taxa, and not even all invaders, are harmful invaders – the last-mentioned should

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plant invasions and invasibility of plant communities 333

Fig. 13.1 Total number of alien plant species, percentage of alien plant species, andnumber of alien plant species per log(area) along the Pacific coast of Americas. ‘Alienspecies’ here are plants growing in individual areas without cultivation. Not all of them arefully naturalized and even fewer are invasive. Nevertheless, numbers of naturalized andinvasive species are proportional to numbers of ‘alien species’ in this diagram. Primarydata or references are in Kartesz & Meacham (1999) and Vitousek et al. (1997).

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334 marcel rejmánek ET AL.

rather be called exotic weeds or exotic pest plants (Booth et al. 2003; Richardsonet al. 2000a).

Weeds are both native and alien and the alien element in weed floras varies overthe world. Most weedy taxa in southern Europe, Malaya, Mexico or Taiwan are native,whereas most weedy taxa in Australia, the USA, Chile, South Africa, New Zealand,Hawai’i, and many other islands are non-native. There may be inherent differencesin invasibility of different parts of the world. Uneven representation of alien, mostlynaturalized, plant species in regional floras along the Pacific shore of the Americasillustrates this point (Fig. 13.1). These differences are certainly partly due to thehistory of human colonization and trade. Nevertheless, similar patterns can also berecognized on other continents (Rejmánek 1996; Lonsdale 1999). For instance, areaswith mediterranean climates (with the exception of the Mediterranean Basin itself )seem to be more vulnerable and the tropics appear more resistant to plant inva-sions. This should not be generalized, however. Savannas and especially disturbeddeforested areas in the Neotropics are very often dominated by African grasses suchas Hyparrhenia rufa and Melinis minutiflora, while similar tropical habitats in Africaand Asia are dominated by Neotropical woody plants, e.g. Lantana camara andOpuntia spp. The absolute number of alien species, therefore, is not necessarily thebest indicator of ecosystem invasibility, at least at this scale. Undisturbed tropicalforests, however, harbour only a very small number of alien plant species and most ofthem do not spread beyond trails and gaps (Rejmánek 1996). It is probably not theextraordinary species diversity of tropical forests that is important but simply thepresence of fast-growing multilayered vegetation that makes undisturbed tropicalforests resistant to invasions.

At the regional scale enormous differences in presence and abundance of invadersamong different communities (ecosystems) within one area seem to be the rule. Anoverview is now available for Central Europe (Table 13.1). Alien species are concen-trated mostly in vegetation of deforested mesic habitats with frequent disturbance(Pysek et al. 2002a,b). In general, native forests harbour a low number and pro-portion of both archaeophytes (introduced before 1500) and neophytes (introducedlater); alien species are completely missing from many types of natural vegetation,e.g. bogs, natural Picea abies forest, and rare in many natural herbaceous commun-ities. Herbaceous communities of extreme habitats and/or with strong native clonaldominants (Nanocyperion flavescentis, Phragmition, Nardion) seem to be most resistantto invasions of both archaeophytes and neophytes. In general, Californian lowlandcommunities (Fig. 13.2) are more invaded than corresponding communities in Europe.However, there are some important similarities. Open and disturbed communitiesare more invaded, while undisturbed forests are less invaded.

Data from California (Fig. 13.2) suggest that proportions of alien species numbersare reasonably well correlated with their dominance (cover). This is probably attribut-able to a simple sampling effect: with increasing proportion of alien species, there isan increasing chance that one or more of them will dominate the community. Whilethere appears to be a general agreement between the proportion of alien speciesnumbers and their actual importance (cover and biomass), some exceptions are quiteremarkable. While alien species number in Chelidonio-Robinion woodland is certainlynot exceptionally high (Table 13.1), the dominant Robinia pseudoacacia is an alien

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Table 13.1 Numbers of alien species, classified according to the time of introduction intoarchaeophytes and neophytes, in representative plant communities of the Czech Republic onthe phytosociological alliance level. Within each vegetation group, alliances are rankedaccording to decreasing total number of alien species. Data from Pysek et al. (2002a).

Number of Number of % invasivearchaeophytes neophytes among

neophytes

Ruderal vegetationSisymbrion officinalis tall-herb comm. of annuals on 96 106 9.4

nitrogen-rich mineral soilsAegopodion podagrariae nitrophilous fringe comm. 16 76 36.8Arction lappae nitrophilous comm. of dumps and 36 45 31.1

rubbish tipsBalloto-Sambucion shrub comm. of ruderal habitats 18 34 41.2Matricario-Polygonion arenastri comm. of trampled sites 20 20 15.0Potentillion anserinae comm. of salt-rich ruderal 12 20 10.0

habitatsConvolvulo-Agropyrion comm. of field margins and 24 16 31.3

disturbed slopesOnopordion acanthii thermophilous comm. of village 34 8 12.5

dumps and rubbish tips

Weed communities of arable landVeronico-Euphorbion weed comm. of root crops on 47 28 21.4

basic soilsPanico-Setarion weed comm. of root crops on 28 15 40.0

sandy soilsCaucalidion lappulae thermophilous weed comm. on 79 11 0.0

base-rich soilsAphanion weed comm. on acid soils 41 8 12.5Sherardion weed comm. of cereals on medium 47 7 14.3

base-rich soils

GrasslandsArrhenatherion mesic meadows 15 56 25.0Festucion valesiacae narrow-leaved dry grasslands 12 12 0.0Bromion erecti broad-leaved dry grasslands 6 8 0.0Nardion subalpine grasslands 0 1 0.0Helianthemo cani-Festucion pallentis rock-outcrop 2 0 –

vegetation

ForestsAlnion incanae ash-alder alluvial forests 4 15 40.0Carpinion oak-hornbeam forests 6 14 14.3Chelidonio-Robinion plantations of Robinia 5 10 60.0Genisto germanicae-Quercion dry acidophilous oak 1 11 36.4

forestsTilio-Acerion ravine forests 5 8 37.5Luzulo-Fagion acidophilous beech forests 0 4 50.0Quercion pubescenti-petraeae thermophilous oak forests 1 2 0.0Quercion petraeae acidophilous thermophilous 0 2 50.0

oak forestsSalicion albae willow-poplar forests of lowland rivers 0 2 50.0

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Table 13.1 (cont’d )

Number of Number of % invasivearchaeophytes neophytes among

neophytes

Alnion glutinosae alder carrs 0 2 0.0Fagion beech forests 0 1 100.0Betulion pubescentis birch mire forests 0 0 –Piceion excelsae spruce forests 0 0 –

Aquatic and wetland vegetationLemnion minoris macrophyte vegetation of eutrophic 0 3 0.0

and mesotrophic still watersCardamino-Montion forest springs without tufa 0 2 50.0

formationPhragmition reed beds of eutrophic still waters 1 1 0.0Magnocaricion elatae tall-sedge beds 0 1 0.0Nanocyperion flavescentis annual vegetation on 1 0 –

wet sand

tree from North America. On the other hand, there are many alien species in somegrassland communities (Festucion valesiaceae, Bromion erecti), but dominants are exclus-ively native and aliens are rarely invasive.

13.3 Invasibility of plant communities

Can we say anything conclusive about differences in invasibility (vulnerability toinvasions) of particular ecosystems? Analyses of ecosystem invasibility based just onone-point-in-time observations (a posteriori) are usually unsatisfactory (Rejmánek1989). In most of the cases we do not know anything about the quality, quantityand regime of introduction of alien propagules. Nevertheless, available evidenceindicates that only very few non-native species invade successionally advanced plantcommunities (Rejmánek 1989; Meiners et al. 2002). Here, however, the quality ofcommon species pools of introduced alien species – mostly rapidly growing andreproducing r-strategists – is probably an important part of the story. These speciesare mostly not shade-tolerant and many of them are excluded during the first 10or 20 yr of uninterrupted secondary succession (Fig. 13.3), or over longer periodsof primary successions. However, some r-strategists are shade-tolerant. e.g. Alliariapetiolata, Microstegium vimineum and Sapium sebiferum. Such species can invade suc-cessionally advanced plant communities and, therefore, represent a special challengeto managers of protected areas.

Plant communities in mesic environments seem to be more invasible than commun-ities in extreme terrestrial environments (Rejmánek 1989). Apparently xeric environ-ments are not favourable for germination and seedling survival of many introducedspecies (abiotic resistance) and wet terrestrial habitats do not provide resources –mainly light – for invaders because of fast growth and high competitiveness of resident

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Fig. 13.2 Native and invasive species in seven plant communities of the Stebbins ColdCanyon Reserve, North Coast Ranges, California (150–500 m above sea level). Eachcolumn represents a mean from three 100-m2 plots. ‘Relative cover’ of invaders is their coverwith respect to the cumulative vegetation cover in all strata (herbs, shrubs and trees).Comparing means for individual vegetation types, the only significant correlation is betweenpercentage of invasive species and total cover of invasive species (r = 0.75; n = 7; p = 0.05).Rejmánek (unpublished data).

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10

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Fig. 13.3 Effect of time since abandonment on the mean species richness of native andexotic species over 40 yr of old-field succession in New Jersey (Meiners et al. 2002). Declineof the mean percentage of exotic species is even more dramatic, decreasing from 58 to 28%.Total and relative cover of exotic species declines significantly as well, but there can be atemporary increase during the first 5 or 10 yr of succession; see also Rejmánek (1989).

species (biotic resistance). We have to be cautious, however, in the interpretation ofthese patterns. When the ‘right’ species are introduced, even ecosystems that have beenviewed as invasion-resistant for a long time may turn out to be susceptible, for instancethe Mojave and Sonoran deserts are facing recent invasions by Brassica tournefortiiand Pennisetum ciliare. Open water is notoriously known as open to all kinds of exoticaquatic plants. In general, disturbance, nutrient enrichment, slow recovery rate ofresident vegetation, and fragmentation of successionally advanced communities pro-mote plant invasions (Rejmánek 1989; Hobbs & Huenneke 1992; Cadenasso &Pickett 2001). In addition, the increasing CO2 level will probably accelerate invasionsin arid ecosystems (Smith et al. 2000).

A general theory of invasibility was put forward recently by Davis et al. (2000):intermittent resource enrichment (eutrophication) or release (due to disturbance)increases community susceptibility to invasions. Invasions occur if/when this situationcoincides with availability of suitable propagules. The larger the difference betweengross resource supply and resource uptake, the more susceptible the community toinvasion. This was anticipated by Vitousek & Walker (1987) (Fig. 13.4) and expressedmore rigorously by Shea & Chesson (2002). Davis & Pelsor (2001) experimentallymanipulated resources and competition in a herbaceous community, and showed thatfluctuations in resource availability of as little as one week in duration could greatlyenhance plant invasion success (survival and cover of alien plants) up to one yearafter such events.

Experiments on invasibility of different types of ecosystems have been gainingmomentum in recent years (Hector et al. 2001; Fargione et al. 2003). Crawley et al.(1999) and Davis et al. (2000) suggested that there is no necessary relationship betweeninvasibility of a plant community and number of species present in that community.

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plant invasions and invasibility of plant communities 339R

esou

rce

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r de

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Time since disturbance

Disturbance

Supply

Demand

Fig. 13.4 Changes in supply and demand of resources after disturbance in terrestrialecosystems. Resource availability is generally at its maximum shortly after disturbance,although conditions of bare ground can inhibit seedling establishment in some sites.Modified from Vitousek & Walker (1987).

Other studies show that such a relationship exists: positive at the landscape scale(e.g. Stohlgren et al. 1999; Sax 2002) and negative at scales of < 1 m2 (neighbourhoodscales sensu Levine 2000). Wardle (2001) provided a valuable methodological criti-cism of many of these studies. Nevertheless, Kennedy et al. (2002) concluded that inherbaceous communities neighbourhood species richness (within 5–15 cm radius)represents ‘an important line of defence against the spread of invaders’. Hubbell et al.(2001) found that in an undisturbed forest in Panama neighbourhood species rich-ness (within 2.5–50 m radius) had a weak but significantly negative effect on focaltree survival. Is there a generalization emerging from studies on neighbourhoodscales? This would not be surprising as vascular plants are sedentary organisms andactual interactions are occurring among neighbouring individuals.

The experimental studies mentioned above relate the number of resident plantspecies to the number and abundance of alien plant species that establish or becomeinvasive. But, the diversity of organisms at other trophic levels in the receiving envir-onment may well be as important, if not more important, than the number of plantspecies. We can expect that diverse assemblages of mutualists (pollinators, seed dis-persers, microbiota that form symbioses with plant roots) would promote invasibility(Simberloff & Von Holle 1999; Richardson et al. 2000b). Recent experiments byKlironomos (2002) on species from Canadian old-fields and grasslands showed thatnative rare plant species accumulate soil plant pathogens rapidly, while invasive speciesdo not. This result has potentially very important consequences. When introduced

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Propaguleinput of

native taxa1

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ACTUALINVASIONRATE IN

A COMMUNITY1

Post-disturbancerecovery rateof residents

Resident herbivoresand pathogensof exotic taxa

Actual amountof availableresources1

Propaguleinput of

exotic taxa1Long-termcharacter of local

environment*2

Non-specificmutualists

Fig. 13.5 Causal relationships between factors and processes which are assumed responsiblefor invasions of exotic species into plant communities. The most important relationships areindicated by thick arrows. * = Spatial heterogeneity, (micro)climate and long-term regimeof available resources and toxic compounds. 1 = Time scale: days–years. 2 = Time scale:years–centuries. The key components are in boxes.

outside of their native territories, plants are often liberated from their enemies,including soil pathogens. This is a clear advantage that makes natives and aliens, atleast temporarily, different (but see Colautti et al. 2004). At the same time, however,as many mycorrhizal fungi can associate with a broad range of plant taxa, rootbeneficial symbionts are likely to be always available to many alien plants.

A conceptual cause-effect diagram (Fig. 13.5) captures all the fundamental aspectsof the ongoing debate on the issue of invasibility. The fact that both invasibility andspecies diversity of residents is regulated in a similar way by the same set of factors– (micro)climate, spatial heterogeneity, long-term regime of available resources –explains why there are so many reports of positive correlation between numbers ofnative and non-native species when several different communities or areas are com-pared. Fast post-disturbance recovery of residents may be a key factor making thewet tropics more resistant to plant invasions – measured as number of invadingspecies per log(area) (Rejmánek 1996).

However, there is very likely one extra factor that is currently poorly understood: thehistorical and prehistoric degree of exposure of resident taxa to other biota (Fig. 13.5).Is this the reason why islands are more vulnerable and Eurasia least vulnerable toinvasions? Is instability of so many man-made monocultures a result of the ‘lack of

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any significant history of co-evolution with pests and pathogens’ (May 1981)? Actualspecies richness may not be as important as the complexity of assembly history.In addition to mathematical models and computer simulations (Law 1999) relevantexperiments with plant communities will have to be designed. Artificial experimentalplant communities that are so often used for invasibility experiments have a clearadvantage of homogeneous substrata and microclimates. However, assembly pro-cesses here are very short and/or artificially directed via arbitrary species pool selec-tion, weeding, reseeding, etc. The existence of well-established phytosociologicalassociations and the fact that plant species are combined in highly non-randompatterns within their natural communities (Gotelli & McCabe 2002) indicate thathistorical assembly processes cannot be substituted by arbitrary mixtures of species.

Finally, longevity/persistence of resident plants is a distinct component of resist-ance to invasions (Von Holle et al. 2003), especially in forest communities, resultingin ‘biological inertia’, including allelopathic chemicals produced by living or deadresidents.

13.4 Habitat compatibility

Identity of exotic taxa (Fig. 13.5) is important for two reasons. First, they may ormay not survive and reproduce in habitats where they are introduced. Second, theymay or may be not spread and become invasive. Recipient habitat compatibility isusually treated as a necessary condition for all invasions. The match of primary (native)and secondary (adventive) environments of an invading taxon is not always perfectbut usually reasonably close (e.g. Beerling et al. 1995; Rejmánek 2000; Widlechner& Iles 2002). In North America, for example, latitudinal ranges of naturalizedEuropean plant species from the Poaceae and Asteraceae are on average 15–20°narrower than their native ranges in Eurasia and North Africa. These differencesessentially reflect the differences in the position of corresponding isotherms and majorbiomes in Eurasia and North America. Major discrepancies between primary andsecondary ranges have been found for aquatic plants where secondary distributionsare often much less restricted than their primary distributions. Vegetative reproductionof many aquatic species seems to be the most important factor. Obviously, secondaryranges, if already known from other invaded continents, should be employed in anyprediction of habitat compatibility.

As for plants introduced (or considered for introduction) from Europe, severaluseful summaries of their ‘ecological behaviour’ are available. Especially the combina-tion of Ellenberg indicator values (Ellenberg et al. 1992) with Grime’s functionaltypes (strategies) (Grime et al. 1988) can be a powerful tool for predictions of habitatcompatibility of European species. The strength of affiliation with phytosociologicalsyntaxa (section 1.4.2) is well known for almost all European taxa. Environmentalconditions (climate, soil, disturbance, management) of all syntaxa are available andpotential habitat compatibility of taxa can be extracted from the European literat-ure. Knowledge of this ‘phytosociological behaviour’ of taxa allows predictions aboutcompatibility with analogous (vicarious) vegetation types, even if these will not alwaysbe correct.

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‘Open niches’, habitats that can support life forms that are not present in localfloras for historical and/or evolutionary reasons, deserve special attention. Dramaticinvasions have occurred in such habitats, e.g. Ammophila arenaria (a rhizomatous grass)in coastal dunes in California, Lygodium japonicum (a climbing fern) in bottomlandhardwoods from Louisiana to Florida, Pinus spp. and Acacia spp. in South Africanshrubby fynbos, Opuntia spp. (Cactaceae) in East African savannas, Rhizophora mangle(mangrove) in tree-less coastal marshes of Hawai’i, and the tree Cinchona pubescens(Rubiaceae) in mountain shrub communities on Santa Cruz Island, Galápagos. Theexplanation of such invasions is confirmed by experiments showing that the competit-ive inhibition of invaders increases with their functional similarity to resident abundantspecies (Fargione et al. 2003).

13.5 Propagule pressure and residence time

The notion of habitat compatibility includes all factors embraced in the concept ofhabitat. Most effort in assessments of habitat compatibility has been devoted to climaticand substrate compatibility, although it is well known that many other factors influ-ence range limits. Invasions result from an interplay between habitat compatibility andpropagule pressure (Fig. 13.5). This is illustrated by the invasion dynamics of theNew Zealand tree Metrosideros excelsa (Myrtaceae) in South African fynbos (detailsin Richardson & Rejmánek 1998). Multiple regression of the number of Metrosiderossaplings on a potential seed rain index (PSRI) and soil moisture revealed that, in thiscase, both factors are about equally important (Fig. 13.6). This example clearly showsthat classification of habitats or communities into ‘invasible’ and ‘non-invasible’ can-not be absolute in many situations. Habitats that are currently unaffected (or onlyslightly affected) by plant invasions may be deemed resistant to invasion. However, aspopulations of alien plants build up and propagule pressure (Foster 2001) increasesoutside or within such areas, invasions could well start or increase. Another aspect isthe propagule pressure of native species: if propagules of natives are not available, asfor instance on abandoned fields in California, the ‘repairing’ function of ecologicalsuccession (Fig. 13.3) does not work.

Residence time – the time since the introduction of a taxon to a new area – repres-ents another dimension of propagule pressure. As we usually do not know exactlywhen a taxon was introduced, we use a ‘minimum residence time’ based on herbariumspecimens or reliable records. Nevertheless, the number of discrete localities of nat-uralized species is significantly positively correlated with minimum residence time(Fig. 13.7). One trivial but important conclusion is that the earlier an exotic pestplant taxon is discovered, the better is the chance of its eradication.

13.6 What are the attributes of successful invaders?

The identity of introduced species certainly matters (Fig. 13.5). Extrapolations basedon previously documented invasions are fundamental for predictions in invasionecology. With the development of relevant databases, this approach should lead to

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Fig. 13.6 The dependence of the sapling density of Metrosideros excelsa on potential seedrain index (PSRI) and moisture in fynbos of the Western Cape, South Africa. PSRI =SUM(1/di), where di is distance to the i-th mature tree in metres within the radius 300 m.The first ordination axis (below) serves as a surrogate for moisture gradient. Standardizedpartial regression coefficients (st. part. regr. coeff.) of the multiple regression are almostidentical. Therefore, both independent variables – environment and propagule pressure –are equally important in this case. M. Rejmánek & D.M. Richardson (unpublished data).

Seed rain (PSRI)

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b·2 (st. part. regr. coeff. for X2) = +0.39, p < 0.005

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Tota

l num

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log (Y) = 1.41 + 0.0074Xn = 63, R2 = 0.38, p < 0.001

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log (Y) = −0.015 + 0.017Xn = 111, R2 = 0.36, p < 0.001

1000

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Minimum residence time (yr)150 200 250

Fig. 13.7 The dependence of the total number of reported localities on the minimumresidence time (yr since the first record) of selected naturalized species in the CzechRepublic and Venezuela. P. Pysek & M. Rejmánek (unpublished data).

immediate rejection of imports of many invasive taxa (prevention) and prioritized con-trol of those that have already been established. Such transregional, taxon-specificextrapolations are very useful in many situations, but our lack of understandingmakes them intellectually unsatisfactory. Understanding how and why certain biolo-gical characters promote invasiveness is very important since even a global databasewill not cover all potentially invasive taxa. In New Zealand, for example, Williamset al. (2001) reported that 20% of the alien weedy species collected for the first timein the second half of the 20th century had never been reported as invasive outsideNew Zealand. For these reasons, several attempts have been made to find differencesin biological characteristics of non-invasive and invasive taxa or, at least, betweennative taxa and non-native invasive taxa in particular floras. Major predictions madeby an emerging theory of plant invasiveness were summarized recently by Boothet al. (2003), Myers & Bazely (2003) and Rejmánek et al. (2004):

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1 Fitness homoeostasis (the ability of an individual or population to maintain relat-ively constant fitness over a range of environments) promotes invasiveness.2 Genetic change can facilitate invasions, but many species have sufficient phenotypicplasticity to exploit new environments.3 Several characters linked to reproduction and dispersal are key indicators ofinvasiveness.4 Seed dispersal by vertebrates is implicated in many plant invasions.5 Low relative growth rate of seedlings and low specific leaf area are good indicatorsof low plant invasiveness in many environments.6 Large native range is an indicator of potential invasiveness. However, severalimportant exceptions are known (Impatiens parviflora, Pinus radiata).7 Vegetative reproduction is responsible for many plant invasions, especially in aquaticand wetland environments.8 Alien taxa are more likely to invade a continental area if native members of thesame genus (and family) are absent, partly because many herbivores and pathogenscannot switch to phylogenetically distant taxa (Agrawal & Kotanen 2003; Mitchell& Power 2003). However, invaders on islands seem to exhibit the opposite tendency(Duncan & Williams 2002).9 The ability to utilize generalist mutualists greatly improves an alien taxon’s chancesof becoming invasive.10 Efficient competitors for limiting resources are likely to be the best invaders innatural and semi-natural ecosystems (Tilman 1999; Shea & Chesson 2002).11 Characters favouring passive dispersal by humans greatly improve an alien planttaxon’s chance of becoming invasive.

Points 3, 4 and 5 are particularly relevant. Reproduction and dispersal are keyissues. Consistent seed production in new environments is usually associated withrather simple or flexible breeding systems. For example, rare and endangered taxa inthe genus Amsinckia (e.g. A. furcata, A. grandiflora) are heterostylic, while derivedinvasive taxa (A. menziesii, A. lycopsoides) are homostylic and self-compatible. Self-pollination has been consistently identified as a mating strategy in colonizing species.Nevertheless, not all sexually reproducing successful invaders are selfers. Pannel &Barrett (1998) examined the benefits of reproductive assurance in selfers versus out-crossers in model metapopulations. Their results suggest that an optimal mating systemfor a sexually reproducing invader in a heterogeneous landscape should include theability to modify selfing rates according to local conditions. In early stages of inva-sions, when populations are small, plants should self to maximize fertility. However,later, when populations are large and pollinators and/or mates are not limiting,outcrossing will be more beneficial, mainly due to increasing genetic polymorphism.

Invasiveness of woody taxa in disturbed landscapes is associated with small seedmass (< 50 mg), a short juvenile period (< 10 yr), and short intervals between largeseed crops (1–4 yr) (see Fig. 13.8 and Rejmánek & Richardson 1996, 2003).These three attributes contribute, directly or indirectly, to higher values of threeparameters which are critical for population expansion: net reproduction rate, recip-rocal of mean age of reproduction, and variance of the marginal dispersal density.For wind-dispersed seeds, the last parameter is negatively related to terminal velocityof seeds which is positively related to √(seed mass) (Rejmánek et al. 2004). Because

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25

16

9

4

1

900

Juve

nile

per

iod

(yr)

576324

14436

0 7 6 5 4 3 2 1

NON-INVASIVE

cemo

cemapal

sab

cou

rig

edu

tor

res

fle

ponnig

ell

mursyl

ban

pin

pat

hal

radcon

str

Seed mass (mg)Interval between

large seed crops (yr)

INVASIVE

lam

K

r

Fig. 13.8 Distribution of 23 frequently cultivated Pinus species in a space created by threebiological variables critical in separating invasive and non-invasive species. The K–r selectioncontinuum running from the upper left to the lower right corner of the diagram alsorepresents the direction of the discriminant function (Z) separating non-invasive andinvasive Pinus species. Z = 23.39 − 0.63√M − 3.88√J − 1.09S, where M = mean seed mass(in mg), J = minimum juvenile period (in yr), and S = mean interval between large seedcrops (in yr). Pine species with positive Z scores are invasive and species with negativeZ scores are non-invasive. Species abbreviations: ban = banksiana; cema = cembra; cemo =cembroides; con = contorta; cou = coulteri; edu = edulis; ell = elliottii; eng = engelmannii;fle = flexilis; hal = halepensis; lam = lambertiana; mur = muricata; nig = nigra; pal = palustris;pat = patula; pin = pinaster; pon = ponderosa; rad = radiata; res = resinosa; sab = sabiniana;str = strobus; syl = sylvestris; tor = torreyana.

of the trade-off between seed number and mean seed mass, small-seeded taxa usuallyproduce more seeds relative to biomass. Invasions of woody species with very smallseeds (< 3 mg), however, are limited to wet and preferably mineral substrata (Rejmánek& Richardson 1996). Based on invasibility experiments with herbaceous species, itseems that somewhat larger seeds (3–10 mg) extend species habitat compatibility(Burke & Grime 1996). As seed mass seems to be positively correlated with habitatshade, large-seeded aliens may be more successful in undisturbed, successionallymore mature plant communities.

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Seed dispersal by vertebrates is responsible for the success of many invaders indisturbed as well as ‘undisturbed’ habitats (Binggeli 1996; Rejmánek 1996; Richardsonet al. 2000b; Widlechner & Iles 2002). Even some very large-seeded alien species likeMangifera indica can be dispersed by large mammals. The proportion of naturalizedplant species dispersed by vertebrates seems to be particularly high in Australia: over50% (Tables 2.1 and 7.3 in Specht & Specht 1999). The assessment of whetherthere is an opportunity for vertebrate dispersal is an important component of thescreening procedure for woody plants (Table 13.2). However, vertebrate seed dis-persal in relation to invasions is complicated (Richardson et al. 2000b).

Table 13.2 General rules for detection of invasiveness of woody seed plants based on valuesof the discriminant function Z*, seed mass values, and presence or absence of opportunitiesfor vertebrate dispersal. Modified from Rejmánek & Richardson (1996).

Opportunities for vertebrate dispersal

Absent Present

Dry fruits and Likely1 Very likely2

seed mass > 3 mgZ > 0 Dry fruits and Likely Likely

seed mass < 3 mg in wet habitats3 in wet habitats3

Fleshy fruits Unlikely4 Very likely5

Not Possibly7

Z < 0 unless dispersedby water6

*Z = 23.39 − 0.63√M − 3.88√J − 1.09S; where M = mean seed mass (mg); J = minimumjuvenile period (yr); S = mean interval between large seed crops (yr). Z was derived on thebasis of a priori defined groups of invasive and non-invasive Pinus species. The function was latersuccessfully applied on other gymnosperms and, as a component of this table, even on woodyangiosperms. Note that parameters in this discriminant function are somewhat different fromthose in Rejmánek & Richardson (1996). This is due to exclusion of Pinus caribaea from the dataset used for estimation of the parameters. This species, that is in general non-invasive in manycountries, is highly invasive in New Caledonia.1 For example Acer platanoides, Cedrela odorata, Clematis vitalba, Cryptomeria japonica, Cytisusscoparius, Pinus radiata, Pittosporum undulatum, Pseudotsuga menziesii, Robinia pseudoacacia, Senna spp.,Tecoma stans.2 Species with large arils (Acacia cyclops) are dispersed by birds.3 For example Alnus glutinosa in New Zealand, Eucalyptus globulus in California, Melaleucaquinquenervia in southern Florida, Tamarix spp. in the south-western USA, Cinchona pubescens inGalápagos and Baccharis halimifolia in Australia.4 Feijoa sellowiana and Nandina domestica are frequently cultivated but non-invasive species in California.The second species, however, is dispersed by birds and water in the south-eastern USA.5 Berberis spp., Clidemia hirta, Crataegus monogyna, Lantana camara, Lonicera spp., Myrica faya,Passiflora spp., Psidium guajava, Rubus spp., Schinus terebinthifolius, Solanum mauritianum.6 Nypa fruticans is spreading along tidal streams in Nigeria and Panama. Thevetia peruviana can bedispersed over short distances by surface run-off in Africa.7 Examples of invasive species in this group are Pinus pinea, Melia azedarach, and Maesopsis eminiiin Africa, Quercus rubra in Europe, Mangifera indica in the Neotropics, and Persea americana inGalápagos.

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Table 13.3 Differences between means of growth related variables for non-invasive,unclassified and invasive Pinus species. Same superscript letters for each variable denotemeans that are not significantly different (p > 0.05, Scheffé test). From Grotkopp et al.(2002).

Variable Non-invasive Unclassified Invasive

n = 8 n = 8 n = 8Relative growth rate (RGR, mg·g−1·d−1) 23a 33b 37b

Net assimilation rate (NAR, mg·cm−2·d−1) 0.505a 0.559a 0.572a

Leaf area ratio (LAR, cm2·g−1plant) 50a 67b 73b

Specific leaf area (SLA, cm2·gleaf) 79a 101b 111b

Relative leaf production rate (leaf·leaf –1·d–1) 0.014a 0.022b 0.024b

Many ecologists assume that a high relative growth rate should be an importantcharacteristic of invasive plant taxa in disturbed or open areas, especially in resource-rich environments. Only a few studies have demonstrated this experimentally (Pattisonet al. 1998; Grotkopp et al. 2002). An analysis of seedling growth rates for 29 Pinusspecies (Table 13.3) revealed that: (i) relative growth rate (RGR) of invasive speciesis significantly higher than that of non-invasive species; (ii) differences in RGR areprimarily determined by leaf area ratio (LAR; leaf area per plant biomass); and (iii)LAR is primarily determined by specific leaf area (SLA; leaf area per leaf biomass).Consequently, invasive species have significantly higher specific leaf area. Moreover,there is a highly significant (R2 = 0.685; p < 0.001) positive relationship betweenRGR and invasiveness of the Pinus species (Fig. 13.8, Table 13.2). Identical resultswere obtained using phylogenetically independent contrasts (Grotkopp et al. 2002).High SLA was implied as an important factor associated with invasiveness of grassesand other plants. In general, SLA < 90 cm2·g−1 (dry leaf mass) seems to be a goodindicator of non-invasive or, at least, less-invasive evergreen woody plants and SLA< 150 cm2·g−1 probably means the same for other vascular plant taxa.

Basic taxonomic units used in plant invasion ecology are usually species or sub-species. However, genera are certainly worth considering as well. Species belongingto genera notorious for their invasiveness or ‘weediness’, e.g. Amaranthus, Echinochloa,Ehrharta, Myriophyllum, should be treated as highly suspicious. On the other hand,a continuum from highly invasive to virtually non-invasive species is also commonin many genera, e.g. Acer, Centaurea, Pinus. Recently, some attention has been paidto taxonomic patterns of invasive plants (Daehler 1998; Pysek 1998; Rejmánek &Richardson 2003). In terms of relative numbers of invasive species, some familiesseem to be over-represented: Amaranthaceae, Brassicaceae, Chenopodiaceae, Fabaceae,Gramineae, Hydrocharitaceae, Papaveraceae, Pinaceae and Polygonaceae. Amongthe larger families, the Orchidaceae is the only under-represented one.

13.7 Impact of invasive plants, justification and prospectsof eradication projects

Numerous studies have documented the wide range of impacts caused by invasiveplants. Many invasive taxa have transformed both the structure and function of eco-

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systems by, for example, changing disturbance- or nutrient-cycling regimes (D’Antonioet al. 1999). In many parts of the world, impacts have clear economic implicationsfor humans, e.g. as a result of reduced stream flow from watersheds in South Africanfynbos following alien tree invasions (Van Wilgen et al. 2001), or through disrup-tion to fishing and navigation after invasion of aquatic plants such as Eichhorniacrassipes.

It is important to stress, however, that the impact of invasive plants on biodiversityis much less dramatic than impact of exotic pathogens, herbivores or predators. Itseems that most of the naturalized/invasive plant species have hardly detectableeffects on biotic communities (Williamson & Fitter 1996; Meiners et al. 2001).There are at least 2000 naturalized plant species in North America and more than1000 of them are invasive. However, not a single native plant species is known tohave been driven to extinction due to interactions with alien plants alone. Even onislands, where numbers of exotic plant species are often increasing exponentially,extinctions of native plant species cannot be attributed to plant invasions per se (Saxet al. 2002). Also, the often reported correlation between numbers of native andexotic plant species on the landscape scale can be interpreted as a lack of mechanismsfor competitive exclusion of native plants by exotic ones. Nevertheless, we should becareful with conclusions – many invasions are quite recent and extinction takes along time.

Considerable progress has recently been made in developing methodologies formaking biological, ecological and economic assessments. In attempting to quantifythe value of ecosystem services of South African fynbos systems and the extent towhich these values are reduced by invasions, Higgins et al. (1997) showed that thecost of clearing alien plants was very small (< 5%) as compared to the value ofservices provided by these ecosystems. Their conclusion was that pro-active manage-ment could increase the value of these ecosystem services by at least 138%. The mostimportant ecosystem service was water, and much work has been done on develop-ing models for assessing the value (in monetary terms) of allocating managementresources to clearing invasive plants from fynbos watersheds. Among the most dan-gerous invaders in riparian areas within the USA are species of the Old World genusTamarix (salt cedar). An instructive economic evaluation of Tamarix impacts isprovided by Zavaleta (2000).

As we showed earlier, the most reliable predictions based on biological charactersare limited to invasiveness (likelihood of species establishment and spread). Predictionsof potential impacts will always be less reliable. Because decline in native speciesrichness is dependent on cover of invaders (Fig. 13.9; Richardson et al. 1989; Meinerset al. 2001), indices based on a ratio of cover to frequency should be tested as impactpredictors for individual taxa. Other obvious impact indicators may be biologicalcharacters of plants that are known to have ecosystem consequences (e.g. high tran-spiration rates or nitrogen fixation).

Invasiveness and impact are not necessarily positively correlated. Some fast-spreadingspecies, like Aira caryophyllea or Cakile edentula, exhibit little (if any) measurableenvironmental or economic impact. On the other hand, some relatively slowly spread-ing species, e.g. Ammophila arenaria or Robinia pseudoacacia, may have far-reachingenvironmental effects (stabilization of coastal dunes in the first case and nitrogen soilenrichment in the second).

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50

40

30

20

10

0

Num

ber

of n

ativ

e sp

ecie

s

0.1 1 10 100 1000

Area (m2)

Fynbos with > 25%cover by aliens

Y = 5.04 + 5.531 log (X)R2 = 0.87(n = 6)

Fynbos without aliens

Y = 15.6 + 10.4 log (X)R2 = 0.595(n = 20)

Fig. 13.9 Species-area relationships for native vascular plant species in South Africanfynbos areas densely infested (squares) by alien woody plants and in uninfested areas(circles). Elevations of the two regression lines are significantly different ( p < 0.001).Sources of the data are acknowledged in Richardson et al. (1989).

There is a need for a universally acceptable, and objectively applicable, term forthe most influential invasive plant taxa within given regions, or globally. A poten-tially useful term to use in this regard is ‘transformer species’ (Richardson et al.2000a). Such species, comprising perhaps only about 10% of invasive species, haveprofound effects on biodiversity and clearly demand a major allocation of resourcesfor containment/control/eradication. Several categories of transformers may bedistinguished:1 Excessive users of resources: water – Tamarix spp., Acacia mearnsii; light – Puerarialobata and many other vines, Heracleum mantegazzianum, Rubus armeniacus; waterand light – Arundo donax; light and oxygen – Salvinia molesta, Eichhornia crassipes;high leaf area ratio, LAR, of many invasive plants (discussed earlier) is an importantprerequisite for excessive transpiration;2 Donors of limiting resources: nitrogen – Acacia spp., Lupinus arboreus, Myrica faya,Robinia pseudoacacia, Salvinia molesta;3 Fire promotors/suppressors: promotors – Bromus tectorum, Melaleuca quinquenervia,Melinis minutiflora; suppressors – Mimosa pigra;4 Sand stabilizers: Ammophila spp., Elymus hymus spp.;5 Erosion promotors: Andropogon virginicus in Hawai’i, Impatiens glandulifera inEurope;6 Colonizers of intertidal mudflats – sediment stabilizers: Spartina spp., Rhizophoraspp.;

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7 Litter accumulators: Centaurea solstitialis, Eucalyptus spp., Lepidium latifolium, Pinusstrobus, Taeniatherum caput-medusae;8 Soil carbon storage modifiers: promotor – Andropogon gayanus; suppressor –Agropyron cristatum;9 Salt accumulators/redistributors: Mesembryanthemum crystallinum, Tamarix spp.

The potentially most important transformers are taxa that add a new function,such as nitrogen fixation, to the invaded ecosystem (Vitousek & Walker 1989).Many impacts, however, are not so obvious. For example, invasive Lonicera andRhamnus change vegetation structure of the forest, affecting nest predation of birds(Schmidt & Whelan 1999), and Lythrum salicaria and Impatiens glandulifera canhave negative impacts on pollination and reproductive success of co-flowering nativeplants (Grabas & Laverty 1999; Chittka & Schürkens 2001).

It follows from the discussion on impacts of exotic plants that careful prioritizationis needed before starting often very expensive and time-consuming eradication projects.Maintenance of biodiversity is dependent on the maintenance of ecological pro-cesses. Our priority should be protection of ecological processes. Attempts to eradi-cate widespread invasive species, especially those that do not have any documentedenvironmental impacts (including suppression of rare native taxa), may be not onlyuseless but also a waste of time and resources. Exotic taxa with large-scale environ-mental impacts (transformers) are usually obvious targets for control and eradication.But when is complete eradication a realistic goal?

There are numerous examples where small infestations of invasive plant specieshave been eradicated. There are also several encouraging examples where widespreadalien animals have been completely eradicated. Can equally widespread and difficultalien plants also be eradicated? On the basis of a unique data set on eradicationattempts by the California Department of Food and Agriculture on 18 species and 53separate infestations targeted for eradication in the period 1972–2000 (Table 13.4),it is shown that professional eradication of exotic weed infestations smaller than 1 hais usually possible. In addition, about 1/3 of infestations between 1 and 100 ha and1/4 of infestations between 101 and 1000 ha have been eradicated. However, costs oferadication projects increase dramatically. With a realistic amount of resources, it isvery unlikely that infestations larger than 1000 ha can be eradicated (Table 13.4).

Table 13.4 Areas of initial gross infestations (at the beginning of eradication projects) ofexotic weeds in California, numbers of eradicated infestations, numbers of on-going projects,and mean eradication effort for five infestation area categories. The data include 18 noxiousweedy species (two aquatic and 16 terrestrial) representing 53 separate infestations. FromRejmánek & Pitcairn (2002).

Initial infestation (ha) < 0.1 0.1–1 1.1–100 101–1000 > 1000No. of eradicated infestations 13 3 5 3 0NO. of on-going projects 2 4 9 10 4Mean eradication effort per infestation (work hours)

Eradicated 63 180 1496 1845 –On-going 174 277 1577 17,194 42,751

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Early detection of the presence of an invasive harmful taxon can make the differ-ence between being able to employ offensive strategies (eradication) and the neces-sity of retreating to a defensive strategy that usually means an infinite financialcommitment. Nevertheless, depending on the potential impact of individual invaders,even infestations larger than 1000 ha should be targeted for eradication effort or, atleast, substantial reduction and containment. If an exotic weed is already widespread,then species-specific biological control may be the only long-term effective methodable to suppress its abundance over large areas (Myers & Bazely 2003).

Regardless of their environmental and/or economical effects, plant invasions pro-vide unique chances to understand some basic ecological processes that are otherwisebeyond the capacity or ethics of standard ecological experiments. We are just begin-ning to fully appreciate these opportunities. However, we have a long way to go toachieve a more complete understanding and more rational decision making.

Acknowledgements

We thank Jennifer Erskine and Jennifer Sedra for comments on a draft of thischapter.

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